COD  0

ln COD _t kt (6.6)

where k is the pseudo-first-order rate constant, which is calculated by plotting

 

 

⁰

ln C OD C OD _t

versus t. The effect of pH on k is shown in Figure 6.3. The value of k increased with increasing pH because the generation of hydroxyl radical was increased at the higher pH [291].

This Figure is ommited due to copyright issue

Figure 6.3. Effect of pH on the pseudo-first-order rate constant (k).

The raw wastewater had a dark-yellow color. On the application of peroxone, decolorization took place rapidly. The variation in pH favored decolorization in the same way as the COD removal. Although the molecular ozone also decolorizes the wastewater effectively despite its selective attacking nature, the color removal rate at the alkaline pH was higher than that in the acidic pH. Figure 6.4 shows the effect of pH on color removal. The effect of the concentration of H2O2 in the peroxone process is well known [292].

This Figure is ommited due to copyright issue

Figure 6.4. Effect of pH on the color removal with respect to time (ozone flow rate: 0.78 mg s^ and [H2O2]: 0.176 mol dm^3).

It was observed that the water was dark yellow initially. After the peroxone treatment, the water started to decolorize rapidly, and after 180 min, the water turned into a clear liquid, as shown in Figure 6.5. Initially, the color disappeared rapidly, and after the final stage of reaction, the removal rate of the color was slower. This behavior was observed probably due to the nature of the products formed during the reaction. For example, during ozonation, initially, aromatic

compounds are converted into phenols and aromatic acids rapidly, then further reaction leads to aldehydes and acids, which have slower reactivity towards ozonation [293].

This Figure is ommited due to copyright issue

Figure 6.5. Decolorization of the pharmaceutical wastewater with respect to time: (a). t = 0, (b). t = 30, (c). t = 60, (d). t = 120, and (e). t = 180 min.

The concentration of the phenolic compounds present in the effluent decreased with time, as shown in Figure 6.6. The degradation of phenolic compounds followed the same trend as the COD and color.

This Figure is ommited due to copyright issue

Figure 6.6. Degradation of phenolic compounds present in wastewater at different pH (ozone flow rate: 0.78 mg s^ and [H2O2]: 0.176 mol dm^3).

The alkaline pH favored the removal due to the presence of the hydroxyl radicals generated in situ. At acidic pH, the removal was found to be 87.5%. At pH 7.5 – 11, phenol was not detected after the treatment. It was observed that the decrease in the removal rate was not commensurate with the COD removal. From Figure 6.6, it is observed that the values of phenol concentration at pH 7.5 and 9 are very close to each other, whereas the slow ozone decomposition at the neutral pH is already known [294]. It has also been reported that the presence of the phenolate ions promotes the generation of hydroxyl radical, which agrees with present behavior [294].

However, the degradation increased with the system pH, and the phenolic compounds were below the detectable range after the treatment.

This Figure is ommited due to copyright issue

Figure 6.7. Removal of ammonia during the peroxone process at different pH (ozone flow rate: 0.78 mg s^ and [H2O2]: 0.176 mol dm^3).

The removal of ammonia by the peroxone process was also studied (see Figure 6.7). The ammonia present in the effluent was converted into nitrate by ozone and hydroxyl radical as follows [295,296]:

3 3 3 2 2

NH  4O  NO

^

 4O  H O H 

^ _(6.7)

3 2 2

N H  H O   N H

^

 H O

^(6.8)

2 3 3 2 2

NH

^

 2O  NO

^

 H O O 

_(6.9)

The removal of ammonia was found to be dependent on pH. Negligible removal of ammonia was observed at acidic pH due to the conversion of ammonia into ammonium, which is recalcitrant towards the oxidants [297]. At pH < 9, less availability of free ammonia and the predominance of ozone as oxidant slow down the removal of ammonia. The generation of bicarbonate ions may increase the removal by suppressing the radical reaction and accelerating the decomposition of ozone [298]. At pH > 9, the hydroxyl radicals take over the removal process and react with free ammonia. Further increase in the pH did not accelerate the removal significantly, although the removal by the peroxone process (>96%) was found to be more effective than ozone alone [297].

The concentration of the chloride ions was found to decrease by the peroxone process, as shown in Figure 6.8.

This Figure is ommited due to copyright issue

Figure 6.8. The concentration of chloride during the peroxone process at different pH (ozone flow rate: 0.78 mg s^ and [H2O2]: 0.176 mol dm^3).

It was mentioned in Section 6.2.2 that molecular ozone dominates as the main oxidant at the acidic pH and the hydroxyl radical plays the role of the dominant oxidant at the alkaline pH [298]. The reactions between the chloride ion and molecular ozone are as follows:

3 2

Cl

^

 O  ClO

^

 O

_(6.10)

3 2 2

ClO

^

 O  ClO  O

_(6.11)

2 3 2 3

ClO

^

 O  ClO

^

 O

^ _(6.12)

2 3 3 3

ClO

^

 O

^

 ClO  O

^ (6.13)

Chloride ions are well known for their property of scavenging the hydroxyl radicals [299]. The reactions between the chloride ion and the hydroxyl radical are as follows [300–302]:

HO Cl  

^

 HOCl

^{ (6.14)}

HOCl

^

 H

^

 Cl   H O

2 ^(6.15)

Cl   OH

^

 HOCl

^{ (6.16)}

Cl   Cl

^

 Cl

2^{ (6.17)}

2 2 2

Cl

^

 Cl

^

 Cl  2Cl

^ _(6.18)

It can be concluded that the gaseous chlorine would be formed at the alkaline pH during the reaction, but the chlorate ions are formed at the acidic pH. The chloride ion removal efficiency was found to have a moderate dependency on the pH.

6.2.3. Effect of H2O2 concentration

The positive effect of the presence of H2O2 on the ozonation process has been well-documented and extensively studied [120,303]. The presence of H2O2 triggers ozone decomposition at a faster rate [223]. Previous studies have suggested that higher H2O2 concentrations may scavenge hydroxyl radicals [120,303]. H2O2 can act as an inhibitor to ozone decomposition by triggering the free radical reactions at elevated concentrations, whereas it accelerates the production of hydroxyl radical at the optimal concentration. Present study also suggests the same behavior inasmuch as the COD removal decreased to 75% for 0.264 mol dm^ of H2O2

whereas, for the H2O2 concentration of 0.176 mol dm^ the removalwas 85%. It is concluded that the COD removal was dropped by 11.8% when the H2O2 dose was increased by 50%.

However, the generation of the superoxide ions may prevent the scavenging of hydroxyl radicals via equation (6.19).

2 2 2 2

H O HO O^H O H ^

(6.19)

The optimal concentration of H2O2 was 0.176 mol dm^because the maximum removal (i.e., 85%) was recorded at this concentrationFigure 6.9 shows the enhancement in the COD removal efficiency with increasing H2O2 concentration.

This Figure is ommited due to copyright issue

Figure 6.9. Effect of concentration of H2O2 on COD removal at pH 7.5 for the ozone flow rate of 0.78 mg s^

For the H2O2 concentrations of 0, 0.088, 0.176, and 0.264 mol dm , the COD removal efficiencies were 61.5, 73.8, 85.4, and 75%, respectively, for same initial pH. Hence, the COD removal efficiency increased with the increasing H2O2 concentration till the optimal H2O2

concentration, which corroborated the results reported [34,38].

This Figure is ommited due to copyright issue

Figure 6.10. Effect of concentration of H2O2 on biodegradability at pH 7.5 for the ozone flow rate of 0.78 mg s^

Biodegradability is considered an essential factor for wastewater treatment. Figure 6.10 shows that the low biodegradability (BOD5/COD = 0.1) of the effluent was enhanced after treatment to 0.55 for the 0.176 mol dm^ H2O2. Several previous studies have also reported that the ozone- based processes generate more biodegradable products on degradation, and hence enhanced biodegradability can be achieved [293,305–308]. Ozone-based treatments reduce aromaticity, resulting in improved biodegradability [309,310]. The presence of H2O2 accelerates the production of hydroxyl radical, thereby generating more biodegradable metabolites than ozone alone due to its non-selective nature. At the optimal H2O2 concentration, the BOD5/COD was increased by 5.5 folds, whereas it increased by three folds at 0.264 mol dm^ H2O2 due to the less availability of the hydroxyl radical. It can be concluded that the peroxone treatment increases biodegradability and the H2O2 concentration plays an important role.

6.2.4. Treatment of the ozonated wastewater by adsorption

After ozonation, the wastewater was subjected to adsorption by GAC.

The a

dsorption on GAC was effective for removing the COD further. The samples obtained from the peroxone process

were directly used for the adsorption studies without any further modification. The initial CODs considerably varied with pH (as shown in Figure 6.11).

This Figure is ommited due to copyright issue

Figure 6.11. Removal of COD during the adsorption on GAC (adsorbent dose: 4 g, temperature: 297 K).

All adsorption experiments were performed at 297 K. During adsorption, the residual COD decreased with time, and after a specific time, the residual COD was constant. It was assumed that the amount of adsorption and desorption of the metabolites were in a dynamic equilibrium at this point. The residual COD and time were termed as equilibrium COD and equilibrium time, respectively.

The kinetic data for the adsorption were analyzed using the pseudo-first-order, pseudo- second-order, and Elovich models at pH 5, 7.5, 9, and 11. The linearized forms of the pseudo- first-order and pseudo-second-order models can be expressed as

 

1

ln q

_e

 q

_t

 ln q

_e

 k t

_(6.20)

2 2

1

t e e

t t

q  k q  q

_(6.21)

where

q

_{e and}

q

_t are the residual COD values at equilibrium and at time t, respectively.

Both of these quantities are expressed in mg g^1. The pseudo-first-order and pseudo-second- order rate constants, i.e.,

k

1 _(min^1_{) and}

k

_{2 (g mg}^1 _min^1), respectively, were obtained from the linear plots of

^ln  q

e

 q

t



versus t and

t q

t versus t, respectively. These plots are shown

in Figures 6.12a and 6.12b. The chemisorption process is best described by the Elovich model [311], as follows:

1 ln

 

1 ln

t e e

e e

q a b t

b b

  _(6.22)

where

a

_{e and}

b

_e are the initial adsorption rate (mg g^ min^) and extent of surface utilization (g mg^), respectively. The values of these two kinetic parameters can be obtained from the plot of

q

_{t versus l n t} (see Figure 6.12c).

Figure 6.12. Adsorption kinetics for the removal of COD by GAC, (a) pseudo-first-order, (b) pseudo-second-order, and (c) Elovich model.

The deviation of the values obtained from the model and the experimental data were calculated in terms of the average relative error (ARE) as follows:

 

^cal _exp^exp

1

ARE % 100 ^{N i i}

i i

q q

N _ q





 ^(6.23)

where

q

cal and

q

exp (mg g^) denote the calculated and experimental adsorption capacities, respectively, and N represents the number of data points. The kinetics parameters calculated from the models are given in Table 6.3.

Table 6.3. Kinetic parameters for the adsorption on GAC (initial COD: 650 mg dm^3, adsorbent dose: 4 g, sample volume: 20 cm³, temperature: 297 K, and contact time: 160 min)

Kinetic Models pH Kinetic Parameters R² ARE (%)

Pseudo-first- order

q

e _(mg

g¹)

k1 (min^1)

5 0.9066 0.0129 0.9965 46.95

7.5 0.9300 0.0076 0.9572 53.80

9 0.9908 0.0058 0.9489 53.79

11 0.9777 0.0053 0.9248 55.92

Pseudo-second- order

q

e _(mg

g¹)

k

2 _{(g mg}^1 min^1)

5 1.4465 0.0222 0.9995 4.64

7.5 1.2387 0.0236 0.9767 7.26

9 1.0973 0.0247 0.9622 13.19

11 1.0436 0.0274 0.9411 17.78

Elovich

a

e_{(mg g}^

min^)

b

e_(g

mg^)

5 0.1039 3.1766 0.9991 0.37

7.5 0.0874 3.8285 0.9345 4.01

9 0.0695 4.2608 0.9303 3.66

0.0869 4.9285 0.8445 6.54

The coefficient of determination

 

^R2 was closer to unity for the fits of the pseudo-second- order model, and the values of ARE were lower too, as compared to the pseudo-first-order model, for all pH. The point of zero charges (pHpzc) of the GAC was in the range of 7.22 – 8 [312–314]. pHpzc plays an important role in the adsorption at different pH. In the present study, the adsorption rate was higher in the acidic medium than in the alkaline medium. At pH <

pHpzc, the surface was positively charged, but at pH > pHpzc, the surface was negatively charged. At pH < pHpzc (i.e., pH = 5), the removal of COD by adsorption was maximum (i.e., 85%). At pH > pHpzc, the adsorption efficiency ranged from 65 – 71%. The good fit of the pseudo-second-order kinetics to the experimental data indicates that the rate-limiting step was chemisorption or ion exchange [149,315,316]. The values of

k

2 ranged from 0.0222 to 0.0274 g mg^1 min^1 for the pH range 5 – 11. At pH 5, the value of

k

2 was minimum, i.e., 0.0222,